Environmental Management (2011) 48:885–894
DOI 10.1007/s00267-011-9745-x
PROFILE
Addressing the Impact of Atmospheric Nitrogen Deposition
on Western European Grasslands
C. J. Stevens • D. J. G. Gowing • K. A. Wotherspoon • D. Alard •
P. A. Aarrestad • A. Bleeker • R. Bobbink • M. Diekmann • N. B. Dise •
C. Duprè • E. Dorland • C. Gaudnik • S. Rotthier • M. B. Soons • E. Corcket
Received: 7 March 2011 / Accepted: 16 August 2011 / Published online: 8 September 2011
Ó Springer Science+Business Media, LLC 2011
Abstract There is a growing evidence base demonstrating that atmospheric nitrogen deposition presents a threat
to biodiversity and ecosystem function in acid grasslands in
Western Europe. Here, we report the findings of a workshop held for European policy makers to assess the perceived importance of reactive nitrogen deposition for
grassland conservation, identify areas for policy development in Europe and assess the potential for managing and
mitigating the impacts of nitrogen deposition. The importance of nitrogen as a pollutant is already recognized in
European legislation, but there is little emphasis in policy
on the evaluation of changes in biodiversity due to nitrogen. We assess the potential value of using typical species,
as defined in the European Union Habitats Directive, for
determining the impact of nitrogen deposition on acid
grasslands. Although some species could potentially be
used as indicators of nitrogen deposition, many of the
typical species do not respond strongly to nitrogen deposition and are unlikely to be useful for identifying impact
on an individual site. We also discuss potential mitigation
measures and novel ways in which emissions from agriculture could be reduced.
C. J. Stevens (&) D. J. G. Gowing K. A. Wotherspoon
Department of Life Sciences, The Open University,
Walton Hall, Milton Keynes MK7 6AA, UK
e-mail: [email protected]
R. Bobbink
B-WARE Research Centre, Radboud University,
PO Box 6558, 6503 GB Nijmegen, The Netherlands
C. J. Stevens
Lancaster Environment Centre, Lancaster University,
Lancaster LA1 4YQ, UK
D. Alard C. Gaudnik E. Corcket
University of Bordeaux, UMR INRA 1202 Biodiversity,
Genes and Communities, Equipe Ecologie des Communautés,
Bâtiment B8, Avenue des Facultés, 33405 Talence, France
P. A. Aarrestad
Norwegian Institute for Nature Research,
PO Box 5685, Sluppen, 7485 Trondheim, Norway
A. Bleeker
Department of Air Quality & Climate Change,
Energy Research Centre of the Netherlands,
PO Box 1, 1755 ZG Petten, The Netherlands
Keywords Acid grasslands Biodiversity Convention
on long-range transboundary air pollution (CLRTAP) Nitrogen deposition Species-rich Nardus grassland
Introduction
The natural global nitrogen (N) cycle has been transformed
by human activities as a consequence of agricultural
M. Diekmann C. Duprè
Institute of Ecology, FB 2, University of Bremen,
Leobener Str., 28359 Bremen, Germany
N. B. Dise
Department of Environmental and Geographical Science,
Manchester Metropolitan University, Manchester M1 5GD, UK
E. Dorland S. Rotthier M. B. Soons
Ecology and Biodiversity Group, Utrecht University,
Padualaan 8, 3584 CH Utrecht, The Netherlands
E. Dorland
KWR Watercycle Research Institute,
PO Box 1072, 3430 BB Nieuwegein, The Netherlands
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intensification and fossil fuel combustion. Between 1860
and 1995 the global creation of reactive N (all forms of N
that are biologically or photochemically active) increased
from 15 to 156 Tg N year-1. Between 1995 and 2005, it
increased by a further 31–187 Tg N year-1 (Galloway and
others 2008). With continued growth of the world population and increasing demand for food, pressures on the
global N cycle are set to increase (Tilman 1999).
Nitrogen oxides are mainly produced by fossil fuel
combustion in transport, power generation and industrial
processes whereas 90% of ammonia emissions come from
agricultural sources (mainly fertilizers and animal manure)
(Erisman and others 2008). Excess reactive N in the
atmosphere is deposited to terrestrial and aquatic ecosystems as wet or dry deposition. Wet deposition occurs when
soluble N compounds are dissolved in rain and cloud
drops; dry deposition consists of gases and particles that
are deposited directly to surfaces. In Europe as a whole,
deposition of N showed a slight decline between 1980 and
2003 (Fagerli and Aas 2008), but in many areas, levels of
deposition remain above those that are known to have an
impact on semi-natural ecosystems. In other parts of the
world, such as developing nations and growing economies,
N deposition is increasing (Galloway and others 2008).
These changes are in contrast to sulfur (S) for which
emissions have been reduced by between 90 and 70% and
deposition has declined rapidly since the 1970s (Fowler
and others 2007).
Atmospheric N deposition can potentially have a wide
range of effects on semi-natural ecosystems including
direct toxicity, increased sensitivity to secondary stress,
acidification and eutrophication (Bobbink and others
2010). Acidification can be caused directly by acid deposition (N and S), but also indirectly through leaching of
basic cations, soil microbial processes and plant uptake.
Increased soil acidity can result in an increased solubility
of metals and a reduced availability of nutrients (Tyler and
Olsson 2001). In Europe, plant species diversity tends to
decline with increasing soil acidity and so species composition and richness can be impacted (Johnston and others
1986; Falkengren-Grerup 1995; Stevens and others 2004).
Because N is the limiting nutrient in many semi-natural
terrestrial ecosystems, the addition of N also has the
potential to increase primary productivity. N deposition
may be less of a threat where phosphorus is the limiting
nutrient. For plant communities the consequence of this
increase in productivity can be a shift towards domination
by species with a high competitive ability under high
resource availability (e.g., Bobbink and others 1998;
Hautier and others 2009).
In acid grasslands (found on soils with a pH of around 5
or below regardless of N deposition status), the main
effects on plant communities are a loss of species richness
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(Duprè and others 2010; Maskell and others 2010; Stevens
and others 2010a), especially forbs, an increase in the
prevalence of competitive species (e.g., Wilson and others
1995), and an increase in the prevalence of acid tolerant
species (e.g., Stevens and others 2010b) resulting in a
species composition which is not typical of this community. These changes are a cause for concern because they
represent a loss of biodiversity across large areas of Western Europe. Changes in plant tissue chemistry have also
been observed on the gradient of deposition (Gidman and
others 2006; Stevens and others 2011b).
Sala and others (2000) identified the five most important
determinants of changes in biodiversity on a global scale:
changes in land-use, increasing atmospheric CO2 concentration, increasing N deposition and acid rain, climate
change, and biotic exchanges (introduction of exotic species). They used a series of scenarios of predicted future
change to identify the relative effects of these drivers on
biodiversity by the year 2100. For global biodiversity,
land-use change was considered the greatest threat followed by climate change and N deposition. The relative
importance of these drivers differed between habitats with
N deposition being among the top three drivers for eight
out of ten terrestrial biomes and the most important in
northern temperate forests.
The Millennium Ecosystem Assessment also identified
N and sulfur deposition together with fertilization as
amongst the most important threats to biodiversity (Mace
and others 2005). It concludes that N inputs are a threat to
biodiversity at the biome and species level.
Both of these global studies are concerned with the
assessment of threat to ecosystems at a global level,
whereas this paper is primarily concerned with a single
vegetation type (acid grassland) in Western Europe.
Semi-natural grasslands are an important component of
European agriculture, supporting extensive grazing and
providing hay. They also support a wide range of plant,
invertebrate and bird species. Acid grasslands are found
throughout Europe in both upland and lowland areas. They
were formerly widespread but, in some areas of Europe,
have been heavily impacted by land-use change and agricultural abandonment (Ellenberg 1996). The habitat we
focus on is identified in the Conservation of Natural Habitats and of Wild Fauna and Flora Directive (92/43/EEC)
(the ‘‘Habitats Directive’’) and is a subgroup of the Natura
2000 habitat ‘‘species-rich Nardus grassland’’.
Focusing on grassland systems in Western Europe,
particularly acid grasslands, this paper aims to assess the
perceived threat of N deposition to biodiversity by policy
makers, identify relevant policy drivers, assess the potential for EU Habitats Directive ‘typical species’ to be used
as indicators of N deposition and identify management
options for mitigating the effects of N deposition. We also
Environmental Management (2011) 48:885–894
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identify outstanding policy-related questions that scientists
and policy makers need to address to effectively protect
biodiversity. To do this we report on the results of a
workshop held for policy makers concerned with N deposition in Western Europe.
Perceived Importance of N Deposition as a Threat
to Biodiversity
In order to assess the perceived threat of N deposition on
biodiversity we asked participants at a workshop for policy
makers and others concerned with environmental policy
(with a focus on national or European conservation and
biodiversity policy) to give their opinions. Workshop participants came from nine countries across Western Europe
(Denmark, France, Germany, Ireland, Netherlands, Norway, Sweden, Switzerland, United Kingdom). The workshop presented results of the European Science Foundation
project ‘BEGIN—Biodiversity of European Grasslands—
Impacts of Nitrogen Deposition’ and was held in October
2009 in Barsac, Gironde, France. The workshop addressed
the following topics: identification of drivers of grassland
habitat change, the use of indicators in assessing impacts of
N deposition on grasslands and mitigation of impacts
through responsive management. In order to assess the
perceived importance of N deposition on grasslands we
asked stakeholders at the workshop to assess the relative
importance of ten drivers of biodiversity loss in grasslands
ranking them from the most important (score 10) to the
least important (score 1). Threats to the biodiversity of acid
grasslands may differ at the Western European scale from
those that we see at a global scale, so the possible drivers
that delegates considered were threats specific to grassland
biodiversity over the next 20 years. The candidate drivers
presented for consideration by the delegates were: atmospheric N deposition, atmospheric sulfur deposition,
intensification of agriculture (including addition of inorganic fertilizers), abandonment of management, fragmentation of habitats, climate change, overgrazing, invasion by
exotic species, recreation and tourism pressures and soil
compaction. There was also an option to add further
drivers.
Abandonment of management was the category that was
most commonly rated as the greatest threat to biodiversity
(45% of delegates), followed by intensification of agriculture and atmospheric N deposition. These were rated as
the top three priorities for many of the delegates as can be
seen by examining the average scores (Fig. 1). Afforestation and building development were both identified as a
threat by one delegate each.
The high score allocated to N deposition reflects the
perceived threat that N deposition presents to biodiversity.
Fig. 1 Average scores allocated by 18 delegates from Western
Europe to identify the most important drivers of grassland biodiversity over the next 20 years. A high score was allocated to the greatest
threats
It also reflects the awareness policy makers have of the
problems that N deposition presents. However, that these
delegates participated in the workshop already indicated
that they were concerned about this issue and the impacts it
may be having in their countries. The results also reflect the
transboundary and widespread nature of the threat presented by N deposition (Fagerli and Aas 2008). Differences
in allocation of scores reflect personal opinion, but also
national priorities and policy. In a country such as the
Netherlands, where N deposition is well publicized and
levels of deposition are high but declining as a result of
successful introduction of abatement measures (Netherlands Environmental Assessment Agency 2005), we saw a
higher ranking than in a country like Ireland, where N
deposition is relatively low. In other parts of the world the
perceived importance of N deposition as a driver of species
change will vary depending on awareness and the perceived importance of other issues. For example, in Eastern
Europe N deposition is high (Dentener and others 2006)
and is thought to be impacting vegetation community
composition (Hejcman and others 2009) but issues associated with land use change and abandonment are more
likely to be considered of high priority (Sikor 2003).
Nitrogen Deposition in European Policy
The transboundary nature of N deposition and other air
pollutants means that European legislation and international conventions are very important in dealing N emission and its impact on biodiversity. Key policy related to
air pollution includes the Convention on long-range
transboundary air pollution (CLRTAP), National Emission
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Ceilings Directive (2001/81/EC), Large Combustion Plant
Directive (2000/80/EC), Ambient Air Quality Assessment
and Management Directive (96/62/EC) and Integrated
Pollution Prevention and Control Directive (91/61/EC and
2008/1/EC). However, although all the above directives are
concerned with reducing air pollution and setting limits for
N emissions, concentrations or deposition rates, many of
these pieces of legislation do not directly consider the
impact of the pollutants on semi-natural habitats. The
CLRTAP is the main exception to this generalization,
considering impacts on human health and environment.
Indeed, of the seven International Co-operative Programmes (ICPs) reporting to the Working Group on Effects
for the CLRTAP, five consider air pollution impacts on
biodiversity.
The 1999 Gothenburg Protocol to abate acidification,
eutrophication and ground-level ozone sets out an obligation to reduce emissions of sulfur, N and volatile organic
compounds and ambient concentrations of ozone below
specified levels. Parties are required to monitor emissions,
ambient concentrations and deposition and collect information on the effects of these pollutants on human health,
terrestrial and aquatic ecosystems and materials. The
Gothenburg Protocol also sets out critical loads and levels
for air pollutants. A critical load is defined as ‘‘a quantitative estimate of an exposure to one or more pollutants
below which significant harmful effects on specified sensitive elements of the environment do not occur, according
to present knowledge’’ (ICP Modelling and Mapping 2004;
Nilsson and Grennfelt 1988). Empirical critical loads for N
deposition are set for a range of habitats (Achermann and
Bobbink 2003) and are based on expert knowledge and
research. They consider the impacts of N on indicators of
biodiversity, such as an increase in the dominance of
N-favored species, decreases in diversity and changes in
soil chemistry. Empirical critical loads are regularly
revised to take account of the most up-to-date scientific
knowledge. Exceedance of these critical loads can be
mapped and provide a tool for determining the potential
damage of N deposition.
Biodiversity policy is also important for assessing the
impact of N deposition. The UN Convention on Biological
Diversity (CBD) and the Ramsar Convention both have the
potential to protect habitats from the deleterious effects of
N deposition (see Bleeker and others 2011) and criticalload exceedance for N deposition is used as one of the
indicators in ‘Streamlining European 2010 Biodiversity
Indicators’ (EEA 2007). Relevant European legislation
includes the Directive on the Conservation of Wild Birds
(2009/147/ES) and on the Habitats Directive. The Habitats
Directive requires member states to take measures to
ensure habitats and wild species are in a favorable conservation status.
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Conservation Management Tools and Options
Detecting N Deposition Impacts Using
‘‘Typical Species’’ as Indicators
In order to fulfill the requirements of the Habitats Directive,
member states are required to undertake surveillance of the
conservation status of the natural habitats and species and to
produce a report every six years. A number of methods are
used to define ‘favorable status’ for each habitat, including
habitat range and area, and presence of typical species.
Typical species are those which can be considered good
indicators of favorable habitat quality, are sensitive to
changes in the condition of the habitat and are detectable by
non-destructive means (European Commission 2006).
Typical species are defined by each member state and are
species considered typical of the habitat. If these species
were suitable as indicators of N deposition impact they
would be very useful for conservation managers.
We undertook an analysis of acid grassland survey data
from Western Europe to determine whether Habitats
Directive typical species are suitable for use as indicators
of N deposition. Typical species were taken from the
Habitats Directive Article 17 database (EEA 2008). The
most appropriate habitat classification under the Habitats
Directive for the grasslands surveyed is ‘species-rich
Nardus grassland’ (habitat code 6230). Only four countries
submitted lists of typical species to the European Environment Agency for species-rich Nardus grassland in the
Atlantic region: Netherlands, Ireland, France and Germany.
This gave a total of 55 typical species with some species
reported for more than one country (Table 1).
The data used in this analysis are from 153 species-rich
Nardus grasslands in ten countries (Belgium, Denmark,
France, Germany, Great Britain, Ireland, Isle of Man,
Netherlands, Norway and Sweden) (Stevens and others
2010a). The grasslands surveyed were selected to cover the
range of atmospheric N deposition in Western Europe and
to give a good range of sites at different latitudes and
longitudes for different deposition values. The grasslands
surveyed all belonged to the association Violion caninae
grassland. The definition of species-rich Nardus grassland
is slightly broader than the definition of the association
Violion caninae grassland (Schwickerath 1944), but Violion caninae can be considered a sub-type of the speciesrich Nardus grassland (Galvánek and Janák 2008; Krahulec
1985). Canonical correspondence analysis (CANOCO 4.5;
ter Braak and Smilauer 2002) was used to show the distribution of species in relation to N deposition. In this
ordination, N and sulfur (S) deposition were used as variables in the analysis whilst soil pH, aluminum concentration, base cation concentration, nitrogen content, carbon
content, C:N ratio, latitude, longitude and mean daily
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Table 1 Typical species for dry acid grassland (habitat code 6230) in
the Atlantic region for countries that submitted data to the Habitats
Directive Article 17 database (EEA 2008)
Table 1 continued
Typical species
Country
Typical species
Thymus pulegioides
FR
Veronica officinalis
DE
Viola canina
DE, FR, IE
Viola lactea
FR
Viola riviniana
IE
Country
Achillea millefolium
IE
Agrostis capillaris
IE
Agrostis curtisii
Antennaria dioica
FR
DE
Anthoxanthum odoratum
IE
Arnica montana
DE
Avenula lodunensis
FR
Botrychium lunaria
DE
Carex arenaria
FR
Carex ericetorum
NL
Carex pallescens
DE
Carex panicea
DE
Carex pilulifera
DE, IE
Chamaespartium sagittale
DE
Dactylorhiza viridis
NL
Danthonia decumbens
FR, IE
Dianthus deltoides
FR
Euphrasia stricta
Festuca filiformis (sub-species of F. ovina)
DE
DE, FR
Festuca ovina
IE
Galium saxatile
DE, FR, IE, NL
Gentiana pneumonanthe
DE
Hypericum maculatum
DE, IE
Hypochaeris radicata
DE
Jasione montana
FR
Juncus squarrosus
DE, IE
Lathyrus linifolius (synonym of L. montanus)
DE
Lathyrus montanus
IE
Luzula campestris
DE, FR
Luzula multiflora
IE
Meum athamanticum
DE
Narcissus bulbocodium
FR
Nardus stricta
DE, FR, IE, NL
Pedicularis sylvatica
Platanthera bifolia
DE, FR, IE, NL
DE, NL
Polygala serpyllifolia
DE, FR, IE, NL
Polygala vulgaris
DE, IE
Potentilla erecta
DE, FR
Pseudarrhenatherum longifolium
FR
Pseudorchis albida
IE
Rhytidiadelphus squarrosus
IE
Rumex acetosella
FR
Sedum anglicum
FR
Serapias lingua
FR
Spiranthes spiralis
NL
Stachys officinalis
NL
Succisa pratensis
IE
Species in bold occurred within more than 5% of the acid grasslands
surveyed
maximum temperature were used as co-variables. There
were no strong geographical trends in the data (Stevens and
others 2011a). For all of the sites, N deposition was
modeled using the best available deposition model.
National models were used for Germany (Gauger and
others 2002), the Netherlands (Asman and van Jaarsveld
2002; Van Jaarsveld 1995, 2004) and Great Britain
(NEGTAP 2001; Smith and others 2000). For all other
countries, the EMEP-based IDEM model (Pieterse and
others 2007) was used.
Many typical species did not occur in sufficient numbers
(more than 5% of sites) in our database to be included in
this ordination analysis. There are two potential reasons for
this. The first is that if a species is particularly common on
a regional level, but not at a broader level, they may make
an ideal typical species for a country, but our dataset may
not have sufficient sites within that region for it to appear in
our dataset at more than 5% of all sites. The second
potential reason that a typical species may not be found in
our dataset is that we have not covered the full range of
types of species-rich Nardus grassland as described in
Natura 2000. Nevertheless, these data can give us an
indication of how suitable a species may be for assessing
the impact of N deposition.
Figure 2 shows the position of the 55 selected Article 17
typical species within an ordination diagram of a canonical
correspondence analysis created using data gathered in the
above-mentioned European survey. Species to the right of
the ordination diagram at the top of the arrow are more
commonly found at high N deposition (based on their
occurrence and cover) within this dataset, those in the
center are neutral with regards to N deposition and those to
the left are more commonly found at lower N deposition. It
is those species on the left that are likely to be most suitable as indicators of low nitrogen deposition impact. As
can be seen from Fig. 2, typical species are scattered across
the ordination diagram, so using typical species lists from
the Netherlands, Ireland, France and Germany, this group
of species is not suitable as an indicator of N deposition
impact in this habitat. Analysis of each of these countries
individually and comparison with national lists showed a
similar scatter of typical species with respect to pollutant
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Fig. 2 Ordination diagram of a canonical correspondence analysis
showing species from 153 Violion caninae grasslands in the Atlantic
biogeographic region of Europe. Filled circles show Habitats
Directive typical species for species-rich Nardus grassland and empty
circles show other species. The empty circles have not been
individually labelled for clarity; a full discussion of species associated
with high and low levels of N deposition is given in Stevens and
others (2011a). The arrows represent increasing total nitrogen
deposition rate (N) and increasing total sulfur deposition rate (S).
Eigen values for axis 1 and 2 are 0.097 and 0.276 respectively. The
total inertia of the analysis was 3.79
deposition (data not shown). There are some individual
species that may be suitable, most notably Agrostis curtisii
(typical species in France), Stachys officinalis (typical
species in the Netherlands) and Polygala serpyllifolia
(typical species in Ireland and Germany). All of these
species are relatively common across Western Europe
(although Agrostis curtisii has a restricted distribution in
the western warm oceanic regions) and so merit further
investigation of their potential as indicators of N deposition
in this particular region.
The presence or absence and abundance of most typical
species are not suitable indicators of N deposition at a
Western European scale, indeed, many of the species
selected as typical species are too common to be useful
indicators even at a national scale. Given the widespread
nature of this result in Western Europe it is likely that this
will be true throughout the whole of Europe. Alternative
species may provide more suitable indicators although
other measures such as Ellenberg N scores, species richness or grass:forb ratio may be more useful, particularly if
temporal changes in these measures could be assessed
(Duprè and others 2010; Stevens and others 2009).
Mitigation of Acidification and Eutrophication
Detecting the potential impacts of atmospheric N deposition is only the initial stage of dealing with the problem.
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Once affected sites are identified, appropriate management
should be put in place to mitigate the effects of deposition.
There are a number of different options for the management of grasslands to control the impact of N deposition.
These fall into two categories, measures to mitigate acidification and measures to mitigate eutrophication.
The main method of mitigating soil acidification is
liming. Liming has a long history of use in agricultural
sciences, indeed the Park Grass Experiment at Rothamsted
Experimental Station in Hertfordshire, England commenced experimental lime addition in 1903 (Silvertown
and others 1994). Liming has been widely used to combat
acidification from atmospheric pollutants and nutrient
addition in many habitats, including grasslands (e.g., Blake
and others 1999; De Graaf and others 1998). Lime (usually
calcium carbonate) reduces soil acidity by exchange of
calcium or magnesium ions with hydrogen ions on soil
particles resulting in a higher soil pH. In a heathlandcatchment liming experiment, Dorland and others (2005b)
found liming resulted in higher soil pH, higher concentrations of base cations and a reduced Al:Ca ratio. Despite
this, there was only a small positive response by vegetation, but changes in species composition were observed
including the cessation of vigorous growth of some competitive species and an increase in rarer species. The
addition of too much lime can damage plants and the
effects if liming can vary depending on the site history, soil
type, amount of N in the soil and the plant species present.
In areas fed by base-rich groundwater, raising water-tables
in the soil has also been used to combat acidification
(Roelofs and others 1996).
Where species composition is changing as a result of
eutrophication and consequent increases in productivity,
measures can be taken to reduce competition for light and
to remove nitrogen from the system. These could include
an increase in cutting frequency or grazing intensity. These
methods are commonly used in the restoration of grassland
from former agricultural land (Walker and others 2003)
and can change species composition, reducing the cover of
productive grasses promoted by the addition of N. Additional winter grazing to disturb the grassland turf is also an
option (Jones and Hayes 1999). The removal of biomass by
cutting and taking off the hay also removes N from the
system with the potential to reduce nutrient status of the
soil and reduce productivity in the long-term. However, the
rate of nutrient removal by this method is usually low
(Hejcman and others 2010), and so it may take some time
for a significant change to be detected (Olff and Bakker
1991). Burning also provides a means of biomass removal
and is traditionally used as a management tool in some
habitats (e.g., heathlands) but it is not commonly used in
grasslands and may not be well accepted by the general
public, because it produces greenhouse gases, particulate
Environmental Management (2011) 48:885–894
pollution and other air pollutants, leaves large areas looking unsightly and generates safety concerns. Another
method that has been tested experimentally for the removal
of reactive N from soils is the addition of carbon (C). C
addition in the form of sucrose, sawdust, starch or cellulose
increases the C:N ratio of soil and can induce microbial
communities to immobilize N in the soil, thus making it
inaccessible to plants (e.g., Eschen and others 2007; Török
and others 2000).
Turf stripping or cutting is the most dramatic method for
the removal of nutrients, but it also removes acidified surface
soil, and so can address both acidification and eutrophication
problems. Turf stripping has been used extensively in some
countries, especially the Netherlands, for the restoration of
heathland (De Graaf and others 1998; Dorland and others
2005a). Topsoil removal is an alternative method that has
been used for the restoration of grasslands (e.g., Buisson and
others 2006) but it presents similar problems to turf stripping. Although turf stripping has undergone experimental
trials in grasslands (e.g., Jansen and Roelofs 1996; Pywell
and others 2002), it is an expensive form of management and
in addition to removing nutrients and acidified soil, it also
removes the soil seed bank and organic matter, as well as
reducing the water holding capacity of the soil (van den Berg
and others 2003a). The removal of the soil seed bank means
that if the local species pool is already depleted, appropriate
species may not be able to re-colonize and may need to be reintroduced (Dorland and others 2004; van den Berg and
others 2003b). Dispersal into large cleared areas is usually
not sufficient for rapid colonization by target species, and
would require large source populations nearby. For perennial species that do not produce large annual seed crops,
unassisted dispersal is likely to be quite limited (e.g., Soons
and others 2005). To facilitate re-colonization of cleared
areas, hay from target communities can be spread out to
supply seeds and improve micro-environmental conditions.
This has been successful in some situations (Poschlod and
Biewer 2005; Coiffait-Gombault and others 2010), but is
limited by the supply of hay from nearby source sites.
Many of these measures are unsuitable for application at
a landscape scale and it would only be appropriate to apply
them to sites of conservation importance or other targeted
areas. Currently it falls to landowners to mitigate against
the effects of N deposition. When conservation organizations are responsible for land management, such mitigation
may be possible albeit expensive, but in many cases the
land is owned by private individuals, so appropriate management needs to be promoted through agri-environment
schemes. Mitigation measures to reduce the effect of N
deposition are not currently incorporated into these
schemes in many parts of Europe, but some of the measures
described above are feasible at different scales. Farmers
need incentives to encourage appropriate management.
891
Payment based on results of management, such as an
increase in species richness or reduction in eutrophic species, could increase motivation and help farmers to value
their land as a conservation resource but may reduce participation in the scheme if farmers are concerned about the
probability of success.
Emission Management Options
Managing the effects of N deposition is frequently
expensive, impractical and in some cases can change the
landscape and the ecosystem dramatically. The only truly
effective and sustainable method to reduce the impact of N
is through the reduction of N emissions.
There are a number of potential methods that could be
used to reduce emissions of N. Emissions of N oxides from
industrial sources are already controlled within Europe to a
large extent through the legislation outlined above. However, this is not always the case world-wide. Options for
the control of reduced N emissions from agriculture have
poor uptake and oxidized N emissions from traffic and
transport are mainly mitigated through efforts to reduce
CO2 emissions.
In principle, the Habitats Directive provides protection
for designated areas of conservation importance. Under the
Directive, projects cannot be approved if they are assessed
to have an adverse effect on a Special Area of Conservation
(SAC) or Special Protection Area (SPA). However, this
decision can only be made if an appropriate review and
assessment is undertaken which, for many agricultural
activities resulting in increased nitrogen deposition to
nearby sites, is not the case. Legislative control of N
emissions from agriculture, and indeed measuring and
monitoring of emissions, provides many challenges due to
the diffuse nature of emissions and the cost of emission
reduction technology. Existing options for reducing N
emissions from agriculture include direct injection of fertilizers and slurries, suitable storage of animal waste, utilizing technologies to minimize fertilizer use and filtering
air before it leaves animal housing areas but all of these
options have the potential to incur additional cost which
may make them unattractive to farmers. Sutton and others
(2011) suggest that improving the nitrogen use efficiency
of crops through improving the genetic potential of crop
varieties, increasing the genetic potential of animals to
increase productivity, improving animal feed quality to
increase feed conversion efficiency and increasing the
efficiency of use of animal manures are key priorities to
reduce agricultural N losses in Europe.
Workshop delegates suggested a number of possible
options for reducing emissions. Diffuse nitrate pollution to
water is currently controlled in Europe through the Nitrates
123
892
Directive (91/676/EEC,) which requires Member States to
identify areas where groundwater nitrate concentrations
exceed 50 mg l-1 or are at risk of doing so. These areas are
designated as Nitrate Vulnerable Zones (NVZs) and Member States must establish ‘Action Programmes’ in order to
reduce and prevent further nitrate contamination. A similar
approach could be taken to N emissions to the atmosphere.
Slurry spreading is an agricultural activity, which is
currently managed for the protection of water quality, but
could additionally be managed for air quality. For example,
regulations in the UK (The Nitrate Pollution Prevention
Regulations 2008) currently prohibit application of farmyard manure to grassland in NVZs between the mid-September and mid-January, when biological uptake is lowest
and runoff is often high. If slurry spreading were also not
permitted during the hottest months of the year, this would
minimize ammonia volatilization.
Reduced meat production and the potential to educate the
public to reduce meat consumption is an effective way to
reduce ammonia emissions. World-wide meat consumption
increased dramatically between 1961 and 1994, and
although the rate of growth in per capita meat consumption
has now slowed in the developed world, this is because it had
already reached a very high level (Rosegrant and others
1999). Using incentives and education to reduce meat consumption could have environmental and health benefits.
Initiatives such as ‘meat free Mondays’ (http://www.
supportmfm.org/), meat free days in schools (such as seen
in Ghent, Belgium), and government recommendations for
reduced consumption of meat (such as seen in Sweden)
are becoming increasingly popular and could reduce N
emissions. Attendees at the workshop signed the ‘Barsac
declaration’ (http://www.nine-esf.org/barsac-declaration)
to reduce meat consumption encourage the availability of
reduced meat portions.
Another suggestion proposed at the workshop was promoting grass-fed animals over housed ones, which generate
higher ammonia emissions. The latter option has benefits
for consumer health (e.g., Daley and others 2010), animal
welfare and the management of the grasslands, since biomass is removed by grazing, but means that emissions
cannot be managed by filtering N compounds from the air.
‘‘Green meat’’ (meat produced on high value grasslands)
provides a further benefit of ensuring that grasslands of
high value for nature conservation remain agriculturally
productive and economically viable. A further option is to
take measures to reduce population growth.
Policy Questions for Research
Results of study and personal observations have convinced
policy makers and scientists from a range of countries and
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Environmental Management (2011) 48:885–894
backgrounds at the workshop that N deposition is having
an impact on species-rich acid grasslands in the Atlantic
biogeographic region of Europe. However, there remain
some important questions that require attention from scientists and policy makers in order for progress to be made
toward providing greater protection for sensitive habitats.
What Changes are of Conservation Concern?
The evidence, briefly summarized above, shows a wide
range of impacts on vegetation and soils, but there is a need
to determine which of these changes are of conservation
concern. For example, it may be that while changes in
chemistry of plant tissues on a gradient of N deposition
have the potential to provide an early warning of plant
stress (Gidman and others 2006), it is likely to be of less
concern in terms of the assessment and management of
sites of conservation importance than changes in community structure and function. A reduction in plant species
richness or a change in community composition may be of
much greater concern, because it represents a loss of biodiversity. Of the changes that are considered of conservation concern in a particular habitat, it is necessary to
determine which of these changes are the most important
and which should trigger the need for action.
How Much Change is Necessary for it to be Considered
Significant?
Once important changes have been identified, the magnitude of these changes needs to be considered. For
example, Stevens and others (2004), in a gradient survey
of acid grasslands in Great Britain, reported an average
reduction in species richness of one species for every
additional 2.5 kg N ha-1 year-1 and suggested that this
pattern was the result of long-term elevated N deposition.
Given a response curve such as this, it is up to scientists
and policy makers to determine ‘acceptable’ changes in
community composition or reductions in species abundance or biodiversity before taking action. The dependence of ecosystem services on species diversity is a
rapidly advancing research front (e.g., Engelhardt and
Ritchie 2001) and will inform the extent to which species
loss can be tolerated.
Addressing these questions will require close collaboration between scientists and policy makers and will provide a future direction for nitrogen-deposition research.
Acknowledgments This project was funded by the European Science Foundation through the EURODIVERSITY-programme, and
national funds were provided by DfG (Germany), NERC (United
Kingdom), NWO (The Netherlands) and INRA, ADEME and Aquitaine Region (France). We are grateful to workshop participants for
their useful contributions.
Environmental Management (2011) 48:885–894
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